Historical trends and ecological risks of polybrominated diphenyl ethers (PBDEs) and alternative halogenated flame retardants (AHFRs) in a mangrove in South China

Science of the Total Environment 599–600 (2017) 181–187

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Historical trends and ecological risks of polybrominated diphenyl ethers (PBDEs) and alternative halogenated flame retardants (AHFRs) in a mangrove in South China Qihang Wu a,b, Xucheng Liu c, Chaozong Liang b, Jonathan Y.S. Leung d,⁎, Huosheng Li a,b, Shejun Chen e, Bixian Mai e, Shenyu Miao f, Yongheng Chen a,b, Zhifeng Wu a,g,⁎⁎, Zhanghe Chen c a

Collaborative Innovation Center of Water Quality Safety and Protection in Pearl River Delta, Guangzhou University, Guangzhou 510006, China Key Laboratory of Water Quality Safety and Protection in Pearl River Delta bMinistry of EducationN, Guangzhou University, Guangzhou 510006, China c College of Life Science, South China Normal University, Guangzhou 510631, China d School of Biological Sciences, The University of Adelaide, Adelaide, South Australia 5005, Australia e State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, China f School of Life Sciences, Guangzhou University, Guangzhou 510006, China g School of Geographical Sciences, Guangzhou University, Guangzhou 510006, China b

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Historical trends and ecological risks of PBDEs and AHFRs are examined. • Concentrations of PBDEs and AHFRs increase since 1997 due to BDE 209 and DBDPE. • DBDPE surpasses BDE 209 as the predominant flame retardant in recent years. • Ecological risk of BDE 209 is high, despite the gradual restriction on production. • Ecological risk of DBDPE appears negligible.

a r t i c l e

i n f o

Article history: Received 4 February 2017 Received in revised form 1 May 2017 Accepted 1 May 2017 Available online xxxx Editor: Adrian Covaci Keywords: Alternative halogenated flame retardant Ecological risk Mangrove

a b s t r a c t While the production of polybrominated diphenyl ethers (PBDEs) was gradually phased out in the last decade, they may still pose hidden danger to the environment due to their toxicity and persistence. On the other hand, alternative halogenated flame retardants (AHFRs) have been increasingly used as substitutes for PBDEs and may further worsen environmental health. To determine the environmental impact of PBDEs and AHFRs, we examined the historical trends and ecological risks of PBDEs and AHFRs in a typical industrialized city in South China by measuring their concentrations in mangrove sediment. Results showed that the concentrations of PBDEs increased abruptly from 1997 to 2009 due to the use of commercial deca-BDE mixture, but were stabilized in recent years. The concentrations of AHFRs, mainly contributed by decabromodiphenyl ethane (DBDPE), kept increasing from 1997 onwards. Based on the temporal trends, DBDPE is predicted to be predominant over BDE 209 in future. Despite the observed similar concentration between BDE 209 and DBDPE, the former posed a

⁎ Corresponding author. ⁎⁎ Corresponding author at: Collaborative Innovation Center of Water Quality Safety and Protection in Pearl River Delta, Guangzhou University, Guangzhou 510006, China. E-mail addresses: [email protected], [email protected] (J.Y.S. Leung), [email protected] (Z. Wu).

http://dx.doi.org/10.1016/j.scitotenv.2017.05.002 0048-9697/© 2017 Elsevier B.V. All rights reserved.

182 Polybrominated diphenyl ether Pollution history

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high ecological risk, while the ecological risk of the latter was negligible. Therefore, more attention is required to manage the contamination of BDE 209 in the environment. © 2017 Elsevier B.V. All rights reserved.

1. Introduction Over the last few decades, polybrominated diphenyl ethers (PBDEs) have been broadly used as flame retardants in a variety of products, especially plastics and electronics (Rahman et al., 2001). Their production and use have been increasing rapidly in many developing countries due to population and economic growth. Despite the effectiveness of PBDEs in reducing flammability, their occurrence in different environmental media has drawn global attention (reviewed in Ma et al., 2012; Law et al., 2014), in view of their toxicity, persistence and bioavailability (Rahman et al., 2001; Darnerud, 2003). In this regard, the production of commercial PBDE mixtures (e.g. penta-BDE and octa-BDE) has been restricted in Europe and North America since the early 2000s (UNEP, 2011). Nevertheless, whether this can effectively regulate the input of PBDEs into the environment remains largely unknown, unless a comprehensive study on the historical trends of PBDEs is conducted. In spite of the restriction, the demand for flame retardants has been annually rising in many developing countries, especially China, in light of the expeditious growth in electronics industry (Mai et al., 2005; Chinn et al., 2014). To meet the more stringent fire safety standards of electronic products, alternative halogenated flame retardants (AHFRs) such as decabromodiphenyl ethane (DBDPE) and dechlorane plus (DP) have been increasingly used as substitutes for PBDEs because their production and use are not prohibited thus far (Covaci et al., 2011). A recent study revealed that the input of AHFRs has been rising rapidly in urbanized and industrialized cities, and will probably surpass PBDEs as the predominant flame retardants in future (Zhen et al., 2016). Since AHFRs have similar chemical structures and properties as PBDEs (Covaci et al., 2011), they are expected to cause similar problems in the environment (e.g. toxicity and persistence). As such, quantifying the concentrations of AHFRs over historical time not only helps evaluate their input in the past, but also manage their use in future. Given the hydrophobicity of PBDEs and AHFRs (Rahman et al., 2001; Covaci et al., 2011), they are mostly adsorbed onto suspended particles in water (e.g. streams and rivers) and end up in the sediment in coastal regions, such as mangroves (Bayen, 2012). Indeed, sediment in coastal regions is regarded as the ultimate sink for persistent organic pollutants (Fu et al., 2003), and thus can act as a good monitoring tool for studying their temporal variation (Wu et al., 2014, 2015). In view of the toxicity and persistence of PBDEs and AHFRs, their accumulation in sediment may pose a long-term ecological risk to the living organisms in mangroves (Environment Canada, 2013). Even worse, the ecological risks of PBDEs may be amplified when the higher brominated congeners (e.g. deca-BDE) are degraded into lower brominated congeners (e.g. tetra- and penta-BDEs) by microbes (Huang et al., 2014), which are more toxic and bioavailable (Darnerud, 2003). Since mangroves provide various vital ecological functions and support a variety of flora and fauna (Ewel et al., 1998; Leung and Cheung, 2017), it is imperative to determine the composition and ecological risks of PBDEs and AHFRs so that their potential impact on mangrove ecosystems can be elucidated. Here, we examined the historical trends of PBDEs and AHFRs in a typical industrialized city in South China by quantifying their concentrations in mangrove sediment along a depth gradient. The composition of PBDEs and AHFRs was also determined to decipher their possible sources over historical time. Since the production of commercial PBDE mixtures was phased out in the early 2000s and the electronics industry in South China started to flourish in the late 1990s, we hypothesized that (1) the concentrations of PBDEs in sediment keep increasing from the late 1990s onwards, but are stabilized or even reduced in recent

years; (2) the concentrations of AHFRs increase more rapidly in recent years. Furthermore, we estimated the ecological risks of PBDEs and AHFRs in mangrove sediment by comparing their concentrations to the guideline values in order to determine whether management effort is needed to regulate their use in future. 2. Materials and methods 2.1. Study site and sampling method A mangrove in Nansha (113°33′00″E, 22°39′14″N), South China, was selected as the study site (average annual temperature: 22.6 °C; average annual precipitation: 1673 mm), which is subject to industrial activities, such as electronic waste (e-waste) recycling, in the region of the Pearl River Delta. An enormous quantity (~50 million mT) of e-waste is annually disposed and processed in this region (Zhang et al., 2009, 2012; Chen et al., 2013), leading to the contamination of PBDEs and AHFRs in the environment. Being located in the distributary of the Pearl River Delta, this mangrove has been influenced by the water flow from the inland cities; therefore, the sediment in this mangrove can act as a good monitoring tool for persistent organic pollutants in South China (Wu et al., 2014, 2015). Sediment samples were collected in December 2011. To assess the pollution history of PBDEs and AHFRs, core sediment samples were collected by a PVC core (10 cm in diameter × 66 cm deep) in the open mudflat of the mangrove (n = 3 replicate core sediment samples). The sediment sample in the core was gradually collected from the top at 2 cm depth interval (i.e. 33 samples per core in total). This interval is based on the previous results of 210Pb dating, where the sedimentation rate is ~ 2 cm yr−1 in the study area (Zhang et al., 2002; Liu et al., 2010). As such, the core sediment sample can represent the pollution history of PBDEs and AHFRs from 1979 to 2011. To estimate the ecological risks of PBDEs and AHFRs, five sampling points were chosen within the mangrove (~100 m between two consecutive points). Then, a rectangular sampler (10 cm long × 10 cm wide × 15 cm deep) was used to collect surface sediment samples (n = 3 replicate surface sediment samples per sampling point). All sediment samples were stored in a freezer at −20 °C prior to analysis. 2.2. Analysis of PBDEs and AHFRs The sediment samples were freeze-dried, ground into powder and passed through a 2 mm sieve. The method used to extract PBDEs and AHFRs was previously described in Wu et al. (2016). Briefly, ~15 g sediment plus small quantity of copper were spiked with surrogate standards (i.e. BDEs 77, 181, 205 and 13C-BDE 209), followed by Soxhlet extraction for 48 h using 200 mL hexane/acetone (1:1, v/v). The extract was concentrated to ~1 mL using a rotary evaporator, followed by purifying in an alumina/silica column. 30 μL internal standard, which contains BDE 118, BDE 128, 4-F-BDE 67 and 3-F-BDE 153, was added into each sample. The concentrations of BDEs 28, 47, 66, 99, 100, 138, 153, 154, 183 (tri- to hepta-BDE congeners), hexabromobenzene (HBB), pentabromoethylbenzene (PBEB), pentabromotoluene (PBT) and 2,3,5,6-tetrabromo-p-xylene (pTBX) were analyzed by Agilent 7890 gas chromatograph equipped with an Agilent 5975 mass spectrometer using electron capture negative ionization (ECNI) in the selective ion monitoring (SIM) mode, separated by a DB-XLB column (30 m × 0.25 mm × 0.25 μm, J&W Scientific). The concentrations of BDEs 196, 197, 202, 203, 206, 207, 208, 209 (octa- to deca-BDE congeners), 1,2-bis(2,4,6-tribromophenoxy)ethane (BTBPE), decabromodiph enyl ethane (DBDPE), polybrominated biphenyl (PBB), anti-dechlorane

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Fig. 1. The mean concentrations of (a) total PBDEs and (b) total AHFRs in sediment along the depth gradient, representing their pollution history from 1979 to 2011.

plus (a-DP), syn-dechlorane plus (s-DP), anti-undecachloropentacyclooctadecadiene (a-Cl11-DP) and anti-decachloropentacyclooctadecadiene (a-Cl10-DP) were analyzed by Shimadzu 2010 gas chromatograph coupled with a QP 2010 mass spectrometer using ECNI in the SIM mode, separated by a DB-5HT column (12.5 m × 0.25 mm × 0.10 μm, J&W Scientific). The GC conditions were described in Tian et al. (2012). Procedural blanks were analyzed once every ten samples to ensure the consistency of instrumental performance. Surrogate recoveries were 86.3% for BDE 77, 80.8% for BDE 181, 96.1% for BDE 205 and 95.3% for 13C-BDE 209 (RSD b 11.0% for all surrogate standards). Total organic carbon (TOC) in sediment was determined using a CHN elemental analyzer (Perkin-Elmer 2400, Perkin Elmer Corp., USA).

Federal Sediment Quality Guidelines (FSeQGs), which are derived from the results of toxicological studies on aquatic organisms (Environment Canada, 2013), and have been widely applied to determine the ecological risks of PBDEs in sediment (e.g. Wang et al., 2015; Liu et al., 2017; Meng et al., 2017). The FSeQGs for tri-BDE, tetra-BDE, penta-BDE, hexa-BDE, octa-BDE and deca-BDE are 44, 39, 0.4, 440, 5600 and 19 ng g−1 (normalized to 1% TOC), respectively. As there is no currently available FSeQG for DBDPE, we adopted 100,000 ng g−1 as the guideline value, below which no toxic effect on aquatic organisms is predicted to emerge (Hardy et al., 2012). This guideline value could conservatively indicate the risk of DBDPE to some common sedimentdwelling organisms (e.g. oligochaetes). The RQ is categorized into three levels: low risk (0.01 ≤ RQ b 0.1), medium risk (0.1 ≤ RQ b 1) and high risk (RQ ≥ 1) (Wang et al., 2015).

2.3. Data analysis 3. Results The ecological risks of PBDEs and AHFRs were estimated by calculating risk quotient (RQ) (Wang et al., 2015), which is a ratio of the concentration measured in sediment to the guideline value. We adopted

The total PBDE concentration in sediment gradually increased from −28 cm to −6 cm, but slightly decreased at the top 4 cm (Fig. 1a, Fig.

Fig. 2. The composition of (a) PBDEs and (b) AHFRs in sediment along the depth gradient.

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Fig. 3. The change in (a) AHFRs/PBDEs and (b) DBDPE/BDE209 in sediment along the depth gradient.

S1 for the concentration of each congener). The total PBDE concentration remained close to 0 between − 28 cm and − 66 cm, indicating that PBDE contamination gradually deteriorated since 1997, but became less severe from 2009 onwards. A similar pattern was observed in total AHFRs, where the concentration was close to 0 between −28 cm and − 66 cm and increased linearly from − 28 cm to the surface (Fig. 1b, Fig. S2 for the concentration of each AHFR). As for the PBDE composition, deca-BDE (i.e. BDE 209) was the predominant congener, contributing N 90% of total PBDE concentration above − 28 cm (Fig. 2a). Its proportion slightly decreased below −30 cm as the relative concentrations of other lower brominated congeners increased. DBDPE was the predominant AHFR above −28 cm, accounting for N 80% of total AHFR concentration (Fig. 2b). The contribution of DBDPE gradually decreased below −28 cm as the relative concentrations of other AHFRs increased, especially DP. The ratio of AHFRs/PBDEs increased from the bottom to the surface sediment, indicating that the relative concentration of AHFRs was gradually increasing over time (Fig. 3a). Similarly, the increasing ratio of DBDPE/BDE209 from the bottom to the surface Table 1 The concentrations of PBDEs and AHFRs in the surface sediment (mean ± S.E.). N.D.: not detected. PBDE

Concentration (ng g−1)

AHFR

Concentration (ng g−1)

BDE 28 BDE 47 BDE 66 BDE 85 BDE 99 BDE 100 BDE 138 BDE 153 BDE 154 BDE 183 BDE 196 BDE 197 BDE 202 BDE 203 BDE 206 BDE 207 BDE 208 BDE 209 ΣPBDEs

0.032 0.093 0.004 0.005 0.058 0.005 0.003 0.062 0.023 0.090 0.518 0.061 0.020 0.071 3.624 0.902 0.188 129.9 135.6

BTBPE DBDPE a-DP a-Cl10-DP a-Cl11-DP s-DP HBB PBEB PBT pTBX

0.061 ± 0.003 142.3 ± 11.4 3.28 ± 0.232 0.107 ± 0.005 0.027 ± 0.002 0.925 ± 0.064 0.014 ± 0.001 0.007 ± 0.000 0.002 ± 0.000 N.D.

ΣAHFRs

146.7 ± 11.7

± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±

0.008 0.022 0.003 0.003 0.023 0.003 0.003 0.026 0.010 0.024 0.518 0.049 0.020 0.071 0.461 0.303 0.188 11.1 12.3

sediment suggests that the use of DBDPE was gradually more extensive than that of BDE 209 since the late 1990s (Fig. 3b). Regarding the surface sediment, BDE 209 was the predominant congener, accounting for ~96% of total PBDE concentration (Table 1). Most of the remaining congeners contributed b 0.1% to total PBDE concentration. DBDPE was the predominant AHFR (~97% of total AHFR concentration), followed by a-DP which only contributed 2.24% to total AHFR concentration. The remaining AHFRs accounted for b 0.1% to total AHFR concentration, except s-DP (0.63%). The RQ of deca-BDE was ~3.2, reflecting the high ecological risk of BDE 209 (Fig. 4). The ecological risks of tri-BDE, tetra-BDEs, hexa-BDEs, octa-BDEs and DBDPE were negligible, while penta-BDEs only posed a low ecological risk as the RQ was below 0.1.

4. Discussion 4.1. Historical trends of PBDEs and AHFRs PBDEs have been the most commonly applied flame retardants in electronic products owing to their effectiveness in reducing

Fig. 4. The risk quotients of PBDEs and DBDPE (mean ± S.E.), indicating their risks to the living organisms in mangrove sediment.

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flammability (Rahman et al., 2001). Given their toxicity and persistence, however, the production of commercial penta- and octa-BDE mixtures was restricted in the early 2000s to minimize the input of PBDEs into the environment (UNEP, 2011). Our findings support the hypothesis that the concentrations of PBDEs started to increase in the late 1990s due to the rapid growth in electronics industry in the Pearl River Delta (Zhang et al., 2009, 2012; Chen et al., 2013). Yet, their concentrations were still increasing until the late 2000s. One of the major reasons is that the use of commercial deca-BDE mixture is still not restricted in most countries, including China (Mai et al., 2005; Ni et al., 2013), explaining the predominance of BDE 209 (~95% total PBDE concentration). The use of commercial deca-BDE mixture is further substantiated by the presence of a small proportion of BDE 206, which matches the commercial formulation of deca-BDE (Saytex 102E, BDE 209: 96.8%; BDE 206: 2.19%) (La Guardia et al., 2006). Saytex 120E is widely used in electrical and electronic equipment, including TV cabinets and cathode-ray tube monitors. Since e-waste recycling industry has been growing fast in South China over the last decade, it explains the gradual increase in the concentration of BDE 209 from 1997 to 2009. Additionally, textile and plastic industries can also contribute to the predominance of BDE 209 due to their rapid growth in South China (Arvanitis and Qiu, 2009). Some previous studies suggested that microbial degradation of BDE 209 can elevate the concentrations of lower brominated congeners (e.g. tetra- and penta-BDEs) in sediment (e.g. Huang et al., 2014; Wang et al., 2015; Yuan et al., 2016), but our findings showed that this process has limited contribution, indicated by the similar proportion of BDE 209 (~95%) from the surface to −26 cm. In view of the elevating demand for flame retardants and global restriction on the production of commercial PBDE mixtures (Covaci et al., 2011), the occurrence of AHFRs in the environment has received more attention in recent years. We revealed that the use of AHFRs started in the same period as PBDEs (i.e. late 1990s), which is probably associated with the growing electronics industry in the Pearl River Delta. The concentrations of AHFRs increased annually, but the increasing rate was augmented in recent years (i.e. the slope from 2004 to 2011 was greater than that from 1997 to 2003, Fig. 1b). This indicates the increasing use of AHFRs and could partially explain why the total PBDE concentration slightly decreased in recent years. Nevertheless, the concentrations of some AHFRs were either negligible (HBB, PBB, PBEB, PBT and pTBX) or only fluctuating (BTBPE) over time, indicating their limited use. We found that DBDPE was the predominant AHFR (~90% total AHFR concentration) since it is commonly used as a substitute for commercial deca-BDE mixture (Covaci et al., 2011). Indeed, the annual production of DBDPE has gradually increased in South China in recent years (Shi et al., 2009; Chen et al., 2013). As a result, DBDPE became more and more dominant over deca-BDE, as reflected by the increasing ratio of DBDPE/BDE209. This finding is corroborated by a number of recent studies. For example, the ratio of DBDPE/BDE209 in sediment ranges from 3.4 to 9.2 in North China (Zhen et al., 2016) and 1.25 in Northeast

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Spain (Barón et al., 2014). Given the gradual phase-out of commercial deca-BDE mixture and increasing annual production of DBDPE, we expect that the concentration of DBDPE will increase substantially in different environmental media in future. Apart from DBDPE, the concentration of DP was also increasing since the late 1990s. DP is heavily produced or imported into the United States (at least 450,000 kg per year), and widely used in coatings for computers and electrical wires (Sverko et al., 2011). In China, the production of DP started in the late 1990s, and therefore the occurrence of DP in the environment was commonly observed in the last decade (Wang et al., 2010). For example, high concentration of DP in sediment was found in Jiaozhou Bay (24.7 pg g−1), Sishili Bay (69.9 pg g−1), Taozi Bay (40.4 pg g−1) and East China Sea (64.4 pg g−1) (Zhao et al., 2011; Wang et al., 2016), which are influenced by industrial activities. To determine the source of DP, the fractional abundance of s-DP (fsyn) can be calculated, where the fsyn value of technical mixture ranges from 0.2 to 0.4 (Wang et al., 2010; Sverko et al., 2011). In this study, the fsyn value ranged from 0.2 to 0.4 from the late 1990s onwards (Fig. S3), indicating the local input of DP probably from industrial processes (e.g. ewaste recycling of DP-containing electronic products). Thus, more attention should be paid to the production and use of DP in future. Contamination of persistent organic pollutants, such as PBDEs and AHFRs, in sediment can be caused by various pathways, such as atmospheric deposition, agricultural runoff and sewage discharge. The low, but detectable, concentrations of PBDEs and AHFRs before 1997 may be due to downward migration through percolation (Zeng et al., 2013). Since then, the abrupt increase in the concentrations of PBDEs and AHFRs implies multiple pathways involved. We suggest that sewage discharge is the major pathway because of the burgeoning ewaste recycling industry and electronics industry in the Pearl River Delta. Owing to the hydrophobicity of PBDEs and AHFRs, they are readily adsorbed onto suspended particles in water, which are transported by water flow and ultimately deposited on sediment. 4.2. Ecological risks of PBDEs and AHFRs Although it is observed worldwide that the concentration of BDE 209 has been stabilized and the concentration of DBDPE is gradually increasing in recent years, substantially overlooked are their ecological risks in the environment. PBDEs can trigger various toxic effects on living organisms (Darnerud, 2003), and thus their accumulation and persistence in sediment can cause a long-term ecological risk to mangrove ecosystems. Despite the similar concentration between BDE 209 and DBDPE, the ecological risk of the former was much higher. This indicates that BDE 209 is still a threat to the environment (see also Yuan et al., 2016), even though its production has been gradually phased out. The ecological risks of PBDEs and AHFRs in other industrialized locations from different parts of the world can be roughly estimated based on their concentrations (Table 2). It is generally observed that BDEs 99

Table 2 The average concentrations of selected PBDEs and AHFRs (ng g−1) in sediment in other locations worldwide. N.D.: not detected. Location

BDE 28

BDE 47

BDE 99

BDE 153

BDE 183

BDE 209

DBDPE

DP

Reference

Nansha mangrove, South China Qi'ao mangrove, South China Dalian Bay, North China Jiaozhou Bay, North China Yellow River Estuary, North China Shanghai River, East China Yangtze River Delta, East China Bui Dau e-waste recycling site, Vietnam Llobregat River Basin, Spain Ebro River Basin, Spain Vltava River, Czech Republic Uhlava River, Czech Republic Durban Bay, South Africa Lenga Estuary, Chile

0.032 0.008 6.73 8.61 / 0.028 / 0.56 / / N.D. 0.29 N.D. /

0.093 0.040 21.5 14.0 0.228 0.344 / 2.2 0.079 0.120 0.47 10.4 174 N.D.

0.058 0.016 40.3 20.6 0.140 0.713 / 3.3 0.166 0.498 0.57 13.64 209 0.79

0.062 0.003 15.5 49.6 0.065 0.045 / 9.8 N.D. 0.146 0.44 1.18 33 /

0.090 0.054 29.0 42.3 0.167 0.065 / 55 0.049 0.486 0.07 2.88 35 N.D.

129.9 33.9 1385 573.6 2.79 13.2 4.57 1700 11.4 8.26 5.35 265.9 3208 0.51

142.3 20.2 4629 1737 / / 6.38 470 11.2 9.68 / / 171 1.54

4.34 0.250 / / / / 1.85 41 0.809 0.516 / / / /

This study Wu et al. (2016) Zhen et al. (2016) Zhen et al. (2016) Yuan et al. (2016) Wang et al. (2015) Zhu et al. (2013) Matsukami et al. (2017) Barón et al. (2014) Barón et al. (2014) Stiborova et al. (2017) Stiborova et al. (2017) La Guardia et al. (2013) Barón et al. (2013)

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and 209 are the major congeners causing high ecological risks. For example, the concentration of BDE 99 in Dalian Bay, Jiaozhou Bay and Durban Bay is far greater than the guideline value; BDE 209 poses a high ecological risk in many places, such as Uhlava River and Bui Dau ewaste recycling site. Alarmingly, the ecological risk of BDE 209 cannot be ignored even in a relatively pristine mangrove in Qi'ao Island, which is located far away from the pollution source in South China. In contrast, DBDPE appears harmless even in the heavily industrialized areas, such as Dalian Bay and Jiaozhou Bay. To date, the toxicity of DP to aquatic organisms remains poorly known, but some studies revealed its low toxicity. For example, it may cause a potential risk to microbes and plants at N 300 μg L−1 (Dou et al., 2015); no acute toxicity was observed in rats even at oral dose of 5000 mg kg−1 (Brock et al., 2010). Given the relatively low DP concentration in sediment (Table 2), the ecological risk of DP is expected to be negligible. More toxicological studies are needed in future to derive the guideline values for other AHFRs so that their ecological risks can be estimated. 5. Conclusion The production and use of commercial penta- and octa-BDE mixtures were globally restricted in the early 2000s. Despite the effectiveness of this regulation in minimizing their input into the environment, we revealed that commercial deca-BDE mixture became heavily used, resulting in severe contamination of BDE 209. While the concentration of BDE 209 has stabilized in recent years due to the gradual phase-out of commercial deca-BDE mixture, it still leads to huge environmental impact. Thus, action should be taken to regulate the input of BDE 209 into the environment (e.g. inspecting the illegal e-waste recycling sites). Regarding AHFRs, we demonstrated the increasing use of DBDPE which will probably surpass BDE 209 as the predominant flame retardant in future. Despite the possible low ecological risk of DBDPE, further studies should be conducted to test the fate of DBDPE, such as degradation in the environment (Wang et al., 2012), and the potential toxicity of its degradation products. Acknowledgements The project is supported by the National Natural Science Foundation of China (41203058 and 41230639), Science and Technology Planning Project of Guangdong Province (2015A020215036 and 2014B030301055), Guangzhou Science and Technology Project (201607010057), and High Level University Construction Project of Guangdong Province (Regional Water Environment Safety and Water Ecological Protection). Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2017.05.002. References Arvanitis, R., Qiu, H., 2009. Research for policy development: Industrial clusters in South China. In: Graham, M., Woo, J. (Eds.), Fuelling Economic Growth. Practical Action Publishing, UK. Barón, E., Gago-Ferrero, P., Gorga, M., Rudolph, I., Mendoza, G., Zapata, A.M., Díaz-Cruz, S., Barra, R., Ocampo-Duque, W., Páez, M., Darbra, R.M., Eljarrat, E., Barceló, D., 2013. Occurrence of hydrophobic organic pollutants (BFRs and UV-filters) in sediments from South America. Chemosphere 92, 309–316. Barón, E., Santín, G., Eljarrat, E., Barceló, D., 2014. Occurrence of classic and emerging halogenated flame retardants in sediment and sludge from Ebro and Llobregat river basins (Spain). J. Hazard. Mater. 265, 288–295. Bayen, S., 2012. Occurrence, bioavailability and toxic effects of trace metals and organic contaminants in mangrove ecosystems: a review. Environ. Int. 48, 84–101. Brock, W.J., Schroeder, R.E., McKnight, C.A., VanSteenhouse, J.L., Nyberg, J.M., 2010. Oral repeat dose and reproductive toxicity of the chlorinated flame retardant dechlorane plus. Int. J. Toxicol. 29, 582–593.

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